SciClaim_sentences_test_withouT.tsv

label	text
2	A newly isolated bacterium was screened out for its survival on medium with 6 000 mg/L kraft lignin as the sole carbon source and energy, and identified as Pandoraea sp. by 16S rRNA gene sequence analysis.
0	The biodegradation experiment was carried out in mineral salt medium, containing 5 000 mg/L kraft lignin as only carbon and energy at pH 8.0 and 30 °C.
2	Under these conditions, significant reduction in color and lignin content by the strain was observed after incubation for 5 d. The strain attains maximum reduction capability in color (44.6%) and lignin content (39.9%) within 5 d of incubation, and reduced chemical oxygen demand(COD) from initial concentration 7 399 to 3 980 mg/L at maximum reduction level of 46.2% on the 4th day.
2	The total ion chromatograph (TIC) of compounds presented in the chloroform extract of control and bacterial treated samples shows the formation of several lignin-related aromatic compounds including some small molecular lignin fragments, indicating a strong destruction in the lignin structure.
0	Lignin is a three-dimensional phenylpropanoid polymer linked mainly by ether or carbon-carbon bonds with an average relative molecular mass over 8 000-11 000.
0	Phenylpropanoid units consist of 4-hydroxyphenyl, guaiacyl, and syringyl structures.
1	The structural complexity of lignin makes it extremely resistant to biological degradation and only a few groups of organisms are capable of degrading lignin polymers, in which fungi are well researched (e.g. white rot fungi).
1	In addition to fungi, bacteria could degrade lignin although having been less extensively studied.
1	Pseudomonas sp., Paenibacillus sp., Aneurinibacillus aneurinilyticus and Bacillus sp., isolated from lignin-containing wastewater, revealed lignin reducing ability.
1	Unlike fungi, it was proposed that bacteria generally degraded lignin in their primary metabolic activity and seldom mineralized lignin, the ligninolytic rates had been found to be lower than those of fungi.
2	In this process, the lignin was converted into p-coumarate, ferulate, vanillate, p-hydroxybenzoate via formation of ""acid-precipitable, polymeric lignin"" intermediates.
1	Compared with fungi, bacteria deserve to be studied for ligninolytic potential because of their immense environmental adaptability and molecular biological versatility.
0	The mechanisms of lignin biodegradation have been the subject of recent research and many simple model compounds including monomeric lignin compounds, such as β-O-4 lignin substructure compounds, lignin-related biphenyl, veratric alcohol and vanillin have been used as substrates.
1	But these treatments could not offer comprehensive evidence to the understanding of bacteria lignin degradation pathways, therefore, high relative molecular mass lignin homogeneous compounds are also needed.
0	Kraft lignin (KL), a by-product in the pulp and paper industry, containing significant similar structure to nature lignin, is a typical lignin model substrate.
0	However, due to the complex structure of KL and the complicated mechanisms of degradation, details about structural modification and the relative molecular mass distributions of KL degradation were not yet clarified.
0	This work was firstly aimed to characterize a novel strain isolated from eroded bamboo slips of the State of Wu dating back to the Three Kingdoms period (220-265A.D.) in China  , which was screened out for its survival on medium with 6 000 mg/L KL as the sole carbon source and energy in our laboratory, then to evaluate its ability to degrade lignin.
0	In order to understand the mechanisms of lignin degradation by the strain, KL was used as a substrate and the detailed chemical modifications induced on KL structure before and after biodegradation were performed using advanced gas chromatography-mass spectrophotometry (GC-MS).
0	The changes of relative molecular mass (M w ) distribution of KL were also investigated by GPC.
0	Meanwhile, various model compounds containing most common linkages found in the macromolecule were also used to understand in depth the metabolic versatility of lignin degradation.
0	The reagents used in this work were of analytical grade.
0	KL powder was purchased from Wuhan East China Chemical Co., Ltd. (Wuhan, China).
0	Monomeric lignin compounds such as vanillin (4-hydroxy-3methoxybenzaldehyde), vanillic acid (4-hydroxy-3methoxy-benzoi acid), trans-ferulic acid (trans-4hydroxy-3-methoxy-cinnamic acid), p-hydroxybenzoic acid, gallic acid (3, 4, 5-trihydroxybenzoic acid), trans-cinnamic acid and guaiacol, were purchased from Sigma (St Louis, Mo, USA).
0	The trimethylsilyl (TMS), a combination of BSTFA (N, O-bis (trimethylsilyl) trifluoroacetamide), trimethylchlorosilane (TMCS) and tetrahydrofurane (THF), were procured from Supelco (Bellefonte, PA, USA).
0	All solutions were prepared in deionized water.
0	A novel bacterial strain, named B-6, was screened out for its survival on medium with 6 000 mg/L KL as the sole carbon source and energy.
0	B-6 was identified by 16S rRNA sequence analysis as genera Pandoraea sp..
0	The exponential growth cells in mineral salt medium (MSM; pH, 8.0; KHPO 4 , 1.0 g/L; K 2 HPO 4 , 1.0 g/L; (NH 4 ) 2 SO 4 , 0.5 g/L; MgSO 4 , 0.05 g/L; CaCl 2 0.01 g/L and trace element solution 1 mL/L) in the presence of 3 000 mg/L KL were harvested by centrifugation at 12 000g for 5 min at 4 °C，then the morphological features were studied by a scanning electron microscopy (SEM).
0	Biochemical tests for bacterial determination were performed using standard methods.
0	The bacterial cell was incubated in MSM liquid medium containing 5 000 mg/L KL at 30°C for 7 d.
0	The culture medium was used as a control in all cases.
0	Samples were taken periodically at 1-day intervals for following analysis: Cells growth was determined by measuring absorbance of inoculated sample at 600 nm (A 600 ) on the spectrophotometer (Shimadzu, UV-2 200) using centrifuged uninoculated medium (named ""control"") as blank.
0	The pH of the samples was measured with pH electrode (Model PHS-3E; Leici, China).
0	For the measurement of residual KL, samples were centrifuged at 12 000g for 5 min. Supernatant (1 mL) was adjusted to pH 8.0 by adding 2 mL NaOH (0.1 mol/L) and absorbance measured at 280 nm on the spectrophotometer.
0	The absorbance value was transformed into lignin mass concentration (mg/L).
0	Then, color in supernatant was determined spectrophotometrically at 465 nm (A 465 ), the absorbance value was transformed into color units (CU).
0	The chemical oxygen demand (COD) was determined by the dichromate method  with modification of incubation temperature from 150 to 165°C as recommended by COD reactor manual (Model 45600-Hach Company, USA).
0	Meanwhile, B-6 was also incubated in MSM medium containing 0.05 mol/L various monomeric lignin compounds (guaiacol, vanillic acid, vanillin, cinnamic acid, p-hydroxybenzoic acid, gallic acid and Ferulic acid) as sole source of carbon and energy, respectively.
0	The plates were incubated at 30 °C for 5 d and each growth was observed.
0	KL biodegradation was monitored by analysis of its metabolites in different incubation time (3 d and 7 d) using GC-MS.
0	Bacterial treated and uninoculated medium samples were centrifuged (12 000g for 5 min) and supernatants were acidified to pH 12 with HCl and then extracted with four volumes of chloroform.
0	The organic layer was collected, then concentrated by rotary evaporator, dewatered over anhydrous Na 2 SO 4 and filtered through 0.22 µm membrane.
0	The extracts were analyzed as trimethyl silyl (TMS) derivatives.
0	For GC-MS analysis, an aliquot of 1µL silylated samples was injected into the GC-MS equipped with Thermo-Finnigan Trace GC Palaris Q, the analytical column connected to the system was a fused silica capillary column DB-5 (30 m×0.25 mm×0.25 mm inner diameter).
0	The GC-MS conditions were as follows: The injector was set at 280 °C in the split mode (1:100 split ratio), oven temperature was held at 80 °C for 5 min and then increased up to 150 °C at 10 °C/min, from 150 °C to 240 °C at 5°C/min and isothermal at 300 °C for 5 min using a heating rate of 10 °C/min in the scan modus.
0	The carrier gas was helium with a controlled flow of 1 mL/min.
0	In the full-scan mode, electron ionization (EI) mass spectra in the range of 30-500 (m/z) were recorded at electron energy of 70 eV. Degradation products were identified by the comparison of retention time and fragmentation pattern, as well as with mass spectra in the NIST spectral library stored in the computer software (version 1.10 beta, Shimadzu) of the GC-MS.
0	The relative molecular mass distributions of the KL and the residual lignin degraded by B-6 were determined using GPC with a HPLC (Waters 244) equipped with Sephadex LH-60 gel columns (pore size 50, 100, 10 3 , 10 4 Å, length 300 mm each, I.D.: 7.8 mm).
0	The acetylated lignin samples were dissolved in tetrahydrofurane (THF) at a concentration of 2 mg/mL and stored for 24 h at 5°C to avoid variations in relative molecular mass.
0	Subsequently, a volume of 20 µL was injected in each sample.
0	THF was used as eluent at a flow rate of 1mL/min at room temperature.
0	A Beckman UV detector was used at 280 nm. For calibration, polystyrene samples with various relative molecular masses were used.
0	The relative molecular masses were calculated from the experimental data using normal GPC calculation procedures.
0	All the experiments were conducted in triplicate under identical conditions and the results were presented as the average value of three data sets.
0	To determine the margin of the error, a confidence interval of 95% was calculated for each set of the samples and the obtained data were analyzed using t-test.
2	Microscopic examination reveals that B-6 cells are short rods with 1.0-1.1μm in length and 0-0.8 μm in width (Fig. 1 ).
2	AY268171.1), corresponding to the 16S rRNA sequence analysis result, biochemical properties of the strain confirms that B-6 is a member of genera Pandoraea sp.
0	Bacterial cells are cultured in MSM containing 5 000 mg/L KL.
2	The changes in color, lignin content and COD during the course of the cells growth in medium, are shown in Fig. 2.
1	A marked increase in cells growth demonstrates that B-6 can utilize KL as its carbon and energy.
2	The color, lignin content and COD are noticeably reduced with the resembled curve trend, which indicates that removal of color and COD reduction is in accordance with the degradation of lignin.
2	During the degradation process, the initial pH of 8.0 decreases to about 7.2 on the second day of incubation then it increases gradually, reaching about 8.4 at the last day (Fig. 3 ).
1	The decrease in pH can be interpreted by the bacterial metabolic activity; bacteria can convert complex lignin into simple organic acids, and it is proposed that the increase in pH is due to simple organic acids which are transferred into other substances as intermediate products.
1	Compared with those reported fungi and bacteria in effectiveness of color and COD reduction, B-6 shows its high lignin degradation activity without other carbon source added in cultivation medium.
2	B-6 attains maximum reduction capability in color (44.6%) and lignin content (39.9%) within 5 d of incubation, with the initial color and lignin content at high 10 551CU and 2 820 mg/L, respectively (Fig. 2 ).
2	Moreover, B-6 reduced COD from initial concentration 7 399 mg/L to 3 980 mg/L at maximum reduction level of 46.2% on the fourth day (Fig. 2 ).
2	RAGHUKUMAR et al  showed that one of marine fungi strains gave 74% decolorization at slightly alkaline pH over a period of 14 d, but high glucose concentration was needed by the fungus to remove color, thus it is economically non-feasible.
2	CHANDRA et al  reported that Bacillus cereus and Serratia marcescens had 45%-62% and 30%-54% reduction in color and lignin content in treatment of pulp paper mill effluent containing KL, with the initial color only at 3 416 CU and lignin content only at 413 mg/L, respectively.
1	Moreover, B-6 can grow on MSM medium containing 0.05 mol/L various monomeric lignin compounds as sole source of carbon and energy under pH of 6.0-8.0 (data not shown).
1	So, it is presumed that B-6 can directly degrade various monomeric lignin compounds and macromolecular KL.
1	Taking all the factors into consideration, B-6 represents a promising candidate for biological treatment of contaminated effluents containing lignin.
1	The ability of B-6 to degrade KL is also confirmed by identifying the low relative molecular mass compounds released in the cultivation medium using GC-MS.
2	The total ion chromatograph (TIC) corresponding to the compounds extracted from cultured supernatants in different incubation time (3 d and 7 d) and its control (uninoculated medium sample) are shown in Figs. 4(a)-(c ) and their identities are listed in Table 1.
2	The compounds detected in the extract degraded by B-6 (Figs. 4(b) and (c )) are different from those of control (Fig. 4(a) ), indicating that an efficient lignin degradation process takes place by B-6.
2	Several aromatic compounds are detected in the extract of medium degraded for 3 d and identified as low relative molecular mass aromatic compounds as well as some small molecular lignin fragments.
2	The peaks at RT (retention time) 32.54 and 36.68 are identified as stilben and biphenyl-type compound, and biphenyl and stilben linkage are the connections between lignin phenylpropane units.
2	Similarly, other low relative molecular mass aromatic compounds are found as veratryl alcohol (RT 11.86), ferulic acid (RT 18.86), 4-hydroxycinnamic acid (RT 21.80) and 4-hydroxy-3-methoxybenzoic acid (RT 29.73).
2	These results are in agreement with the pH (Fig. 3 ) and lignin content decrease findings (Fig. 2 ).
1	The release of these low relative molecular mass lignin-related aromatic compounds has been reported to be lignin degradation intermediate in Phanerochaete chrysosporium as well as in bacteria.
2	The compounds detected in the extract degraded by B-6 for 7 d are different from those for 3 d (Fig. 4 (c) ).
1	Interestingly, the small molecular lignin fragments disappear, which are thought to be further metabolized as metabolic intermediates.
2	Some aldehyde (RT 23.31, RT 27.05), ketone (RT 12.33, RT 20.15, RT 21.85, RT 32.39), Table 1 Compounds identified as TMS derivatives in chloroform extract from control and bacterial degraded kraft lignin samples as given in Fig.
2	Retention time/min ln Fig. quinoid-type (RT 12.33) compounds are detected, whose presences are in accordance with the literature that side-chain cleavage between lignin units led to further form aromatic aldehydes and ketones.
2	It should be noted that the biphenyl-type compound (RT 36.68), which has been reported very stable and its decomposition seems to be the rate-limiting step in lignin degradation  , decreases sharply compared to 3 d, indicating that B-6 has strong capability for a more active degradation of biphenyl compounds compared to other lignin degrading bacteria.
2	Ferulic acid, which is involved in linking the lignin and hemicellulose fraction of lignocellulose, is identified in the extract degraded by B-6.
2	However, vanillic acid and vanillin, metabolic products of ferulic acid, are not detected during the whole incubation time.
2	Further more, there is significant increase in the content of ferulic acid at 7 d compared to 3 d.
2	This is not coincident with the Ref.. But in our present work, B-6 can grow in the medium containing ferulic acid and vanillic acid (approximate pH 6.5, data not shown) as only carbon and energy respectively.
0	The reason may be due to the enzymes involved in degradation of ferulic acid can not occur where the cells growth is in inadaptable condition and the medium was at the alkaline pH.
2	A major factor determining the rate of bacterial lignin degradation is the relative molecular mass, and KL biodegradation induced by B-6 is further studied by evaluating the changes in relative molecular mass distribution shown in Table 2 and Fig.5 using GPC.
1	The results clearly demonstrate that high relative molecular mass lignin fragments are biodegradable by B-6, which is in accordance with the previous results (Fig. 2 and Fig. 4 ).
2	As shown in Table 2 , KL has much high relative molecular mass ranged between 2 770 and 12 000.
2	The degradation of KL is efficient during the first 3 d, the fraction with high relative molecular mass (>12 000) Fig.5 Relative molecular mass distributions of KL degraded by B-6 in different incubation time decreases from 13.1% to 9.1%, while the fraction (6 520<M w <12 000) increases from 19.7% to 22.6%, meanwhile, the fraction with low relative molecular mass (<2 770) begins to increase from 54.6% to 58.9%, the results mean it is probable that the bonds among lignin precursors are cleaved to some small molecular lignin fragments.
2	Compared to 3rd day, the profile of 7th day is shifted to a smaller region and the height around the fraction (6 520< M w <12 000) is lower as shown in Fig.5 , at the same time, great change in the fraction (M w <2 770) is also observed, the strong increase from 58.9% to 76.3% is likely attributable to the fact that the intermediate products are further partly degraded to some smaller lignin aromatic compounds, resulting in an increase of the low relative molecular mass lignin.
1	In conclusion, the metabolic products as GC-MS detected and GPC profiles after biodegradation compared to the control in this work are postulated a possible mechanism.
1	Firstly, the bonds between lignin units are directly cleaved, some low relative molecular mass aromatic compounds and small molecular lignin fragments are produced as metabolic intermediates, then these compounds are oxidized to aromatic aldehydes, ketones and so on, meanwhile, the stable biphenyl-type compounds is also degraded.
0	In the present work, the enzymes secreted by B-6 and the cofactors in culture which are responsible for degrading kraft lignin have not been identified and elucidated.
0	The enzymes involved in the degradation of lignin and various monomeric lignin compounds have been fully reported by other researchers.
1	Laccases lignin peroxidase (Lip) and manganese peroxidase (MnP) are believed to be involved in biodegradation of lignin and of a wide variety of lignin-model compounds.
1	Similarly, it is assumed that of these peroxidases were involved in breaking down kraft lignin.
1	However, a complete characterization of the ligninolytic enzymes and the cofactor in culture requires further studies to determine their definitive role in the degradation of lignin.
1	Although the results from the present work on the degradation of KL by B-6 are cautiously interpreted, lignin degradation mechanisms are rather complicated; The more information about the bacterial degradation should be further explored.
1	And the utility of B-6 could be extended to other biotechnological areas such as the pulp and paper industry, dyeing and printing industry.
1	For instance, the ability of B-6 to secrete involvedimportant extracellular enzymes can be efficient to degrade some refractory macromolecules such as lignin and anthracene.
0	However, further experiments would lead to test the hypotheses. Studies on the formation and expression of ligninolytic enzymes on KL degradation by Pandoraea sp B-6 are in progress in our laboratory.
1	1) A new bacterium, identified as Pandoraea sp.B-6 on the basis of 16S rRNA gene sequence analysis, has efficient ability to degrade KL and lignin-model compounds.
2	B-6 is able to remove 44.6%, 39.9% and 46.2% of color, lignin content and COD within only 5 d of incubation in medium containing 5 000 mg/L KL, repectively.
1	2) The appearances of various intermediates and metabolic products in medium indicat that B-6 is capable of degrading and KL can be utilized as sole source of carbon.
1	Further, B-6 can degrade the lignin intermediate metabolic product-biphenyl compound, which is reported very stable and is the rate-limiting step in lignin degradation.
1	3) The relatively high removal efficiency shows that B-6 has the potential for treatment such as paper mill effluent and printing and dyeing wastewater containing lignin-derived compounds.
0	Decomposition rates for crop residues have generally been estimated based on data obtained using buried litter bags.
0	Because of limited soil-residue contact, the litter bag technique may not adequately simulate decomposition when residues are mixed through the soil.
0	In field microplots, decomposition of wheat (Triticum aestivum) and barley (Hordeum vulgare) straw (autumn-incorporated at a rate of 7 t/ha) mixed through the 0-0.20 m soil layer was compared with straw decomposition in fibreglass bags (4-mm mesh) buried at a depth of 0.20 m. A surface-placed straw treatment and a no-straw control were included for comparison.
2	Emissions of CO 2 were monitored from the incorporated straw treatments and undecomposed straw was recovered after 158 days (mean soil temperature during the trial period was 8 ˚C at 0.10 m).
2	Emissions of CO 2 from the soil-mixed straw treatment were generally greater than from the buried bag treatment in the 2 months following incorporation.
2	Output of CO 2 -C over the first 73 days averaged 83 g/m 2 for the soilmixed straw treatment compared with 61 g/m 2 for the litter bag treatment and 34 g/m 2 for the no-straw control.
1	Over the entire trial, CO 2 -C attributable to straw (CO 2 -C output from straw-treated plots minus CO 2 -C emitted from the control) was 66% greater for soil-mixed straw than for litter bag straw, indicating that within soil placement can have a strong and persistent effect on straw decomposition.
1	Straw type had a small but significant (P < 0.05) effect on CO 2 output (barley > wheat).
2	Straw mass loss during the trial averaged 66% for soil-mixed straw, 32% for litter bag straw, and 13% for straw placed on the soil surface.
1	The low recovery of soil-mixed straw is partly due to difficulty of extracting small (<2 mm) residue fragments from the soil; however, such fragments could legitimately be considered part of the soil organic matter.
1	The results confirm that straw that is well distributed through the soil may decompose more rapidly than would be anticipated from litter bag measurements.
0	Post-harvest crop residues are an important resource for arable farmers because they add organic matter to the soil and recycle plant nutrients.
0	Returning crop residues to the soil has long been advocated as a practice that helps sustain soil quality and fertility over the long-term (Campbell and Zentner 1993).
0	Post-harvest residues may affect several factors that influence yield of the following crop, including plant establishment, nutrient availability, soil temperature, water infiltration and evaporation, and disease and pest populations.
0	The challenge is to manage residues in a way that maximises the benefits while minimising problems associated with their retention.
0	It is especially important that undecomposed residues do not hinder mechanical operations suchasseeddrilling.
0	Athoroughknowledgeofthe factors affecting decomposition is a pre-requisite to develop best management practices for crop residues.
0	Decomposition is a complex process that is influenced by the chemical and physical properties of the residue (e.g. lignin and N contents, particle size) and by soil factors that control microbial activity (Parr and Papendick 1978).
1	Due to more favourable conditions for microbial activity, soil-incorporated residues normally decompose at a faster rate than residues on the soil surface (Parr and Papendick 1978; Christensen 1986; Douglas and Rickman 1992; Curtin et al. 1998; Wang et al. 2001).
1	Interactions between residues and the soil create the environmental conditions under which decomposing microorganisms function.
1	Most importantly, these interactions control the water potential of the straw, the rate of microbial colonisation, and N supply to the decomposing microorganisms (Parr and Papendick 1978; Wang et al. 2001).
1	The disposition of residues within the soil may affect the rate of decomposition.
0	Where residues are incorporated by ploughing, they become concentrated beneath the inverted plough layer and limited soil-residue contact may slow microbial colonisation and water absorption by the residues and restrict N availability to sites of decomposition.
0	When residues are mixed through the soil, as occurs when the soil is cultivated using a disced implement, these constraints are alleviated because of intimate and extensive soil-residue contact.
0	Little work has been done to quantify the effects of withinsoil placement on residue decomposition (Kanal 1995).
1	From a laboratory study, Parr and Reuszer (1959) reported that early decomposition was slower where straw was localised rather than mixed through the soil, but total decomposition over the 6-week incubation differed little between the 2 placement methods.
0	However, caution is required when extrapolating laboratory decomposition data to the field (Parr and Papendick 1978).
0	Residue decomposition may be determined either by measuring mass loss (e.g. Douglas and Rickman 1992; Beare et al. 2002) or by measuring CO 2 -C evolution from decaying residue (e.g. Bremer et al. 1991; Jensen et al. 1997; Berg and McClaugherty 2003).
0	Because of operational simplicity, mass loss using a litter bag technique has been the preferred method in many decomposition studies (Heal et al. 1997).
0	Crop residue, contained in fibreglass mesh bags, is placed in the soil at a depth of 0.15 or 0.20 m and undecomposed residue measured by recovering the litter bag containing the remaining material (Beare et al. 2002).
0	The litter bag technique simulates the placement achieved when residue is incorporated by ploughing.
0	However, the technique has been criticised on the grounds that the mesh could hinder entry of soil animals (earthworms and arthropods), which may be important decomposers in some systems.
0	Possibly a more serious criticism of the litter bag method is that it can underestimate decomposition because of poor soil-residue contact (Cogle et al. 1987).
0	Although soil-mixed residues may give a more meaningful estimate of decomposition than the litter bag technique, recovery of unconfined residues is problematic because decomposition and fragmentation result in residue particles too small to be physically removed.
0	The unrecovered residue could be quantified using 14 C labelling techniques, but a high level of technology is needed to produce and analyse the labelled material (Cogle et al. 1987).
0	Measurement of CO 2 -C output from decomposing residue provides an alternative method of evaluating the effects of within-soil residue placement.
0	Although monitoring of CO 2 emissions during the decomposition period is time-consuming, it does provide information on the dynamics of the decomposition process at time scales that are not possible using the litter bag technique, which, as normally employed, measures mass loss over relatively long time steps (several weeks).
0	Data obtained using the litter bag technique have been used to develop simple models to predict the decomposition of incorporated cereal straw (Douglas and Rickman 1992).
0	These models have been tested against litter bag data from a range of studies (Douglas and Rickman 1992 ), but we do not know if they realistically predict decomposition when residues are distributed through the soil.
0	In this paper we report on a field study comparing decomposition of cereal straw measured using the litter bag technique (simulating the residue placement achieved by ploughing) with decomposition of unconfined residues mixed through the soil (simulating disc incorporation).
0	The experiment was on a Wakanui silt loam (Immature Pallic Soil; Hewitt (1993) ) at the Crop & Food Research farm near Lincoln, Canterbury, New Zealand (43 ˚38 0 S, 172 ˚30 0 E).
2	The sand, silt, and clay contents were 180, 620, and 200 g/kg, respectively.
0	At the start of the experiment, the soil (0-0.20 m) had total C and N contents of 23.2 and 1.7 g/kg, respectively, mineral N content of 47 mg/kg, microbial biomass C, determined by the fumigation-extraction method (Vance et al. 1987; Horwath and Paul 1994) of 207 mg/kg; and pH ~6.
0	Bulk density values for the 0-0.075, 0.075-0.15, and 0.15-0.20 m layers were 1.15, 1.16, and 1.31 Mg/m 3 , respectively.
0	The site had been fallow in the growing season before our experiment.
0	This was preceded by a cereal crop (1999) (2000) and another fallow season (1998-99).
0	Soil moisture content at initiation of the experiment was 14.0% (gravimetric basis).
0	Treatments and trial establishment Wheat (Triticum aestivum L.) and barley (Hordeum vulgare L.) straw used in the experiment was collected as standing stubble from plots on the Crop & Food Research farm in the month preceding the experiment.
0	The initial composition of the straws is given in Table 1.
0	The straws were air-dried and cut into 50-mm lengths before use in the experiment.
0	The trial was laid out as an incomplete Latin square with 5 replicates of 7 treatments.
0	The treatments were: (1) control (no added straw); (2) barley straw (in 4-mm mesh bags) placed horizontally in the soil at a depth of 0.20 m (simulated plough treatment); (3) wheat straw (in 4-mm mesh bags) placed horizontally in the soil at a depth of 0.20 m; (4) barley straw mixed through the 0-0.20 m soil layer (simulated disc treatment); (5) wheat straw mixed through the top 0.20 m of soil; (6) barley straw placed on soil surface; and (7) wheat straw placed on soil surface.
0	The straw application rate was equivalent to 7 t/ha, which is about the average rate of straw production in the Canterbury region.
0	The experiment was established on 6 April 2001 (i.e. early autumn).
0	Soil was dug from microplots (0.30 by 0.30 by 0.20 m deep) by hand.
0	The sides of the holes (but not the bases) were lined with polythene to prevent lateral diffusion of CO 2 from the microplots.
0	In the simulated disc treatments, straw was thoroughly mixed through the excavated soil before repacking it into the hole.
0	In the litter bag treatments, mesh bags (0.30 by 0.30 m; 4-mm mesh) containing straw were placed at the bottom (0.20 m depth) of the holes, and the excavated soil was well mixed before refilling the holes.
0	The control plots and plots to be used for the surface straw treatments were also excavated and the soil mixed before returning it to the holes.
0	The microplots were packed to approximately the original bulk density.
0	To prevent removal by wind, surface-placed straw was covered with mesh (4 mm) that was anchored to the ground using metal pegs.
0	Plots were kept plant-free during the experiment by hand removing seedlings as they emerged.
0	Soil moisture content (0-0.20 m) was measured periodically during the experiment by time domain reflectrometry using 20 cm waveguides.
0	Emissions of CO 2 were monitored from the day following trial establishment and measurements were usually made every 2-3 days until emissions declined due to low soil temperatures in winter, when less frequent monitoring was deemed adequate.
0	A plastic chamber (0.24 m diam.; height 0.25 m) with a sharpened base was pressed into the soil to a depth of 40 mm (the chamber covered 50% of the plot area).
0	The CO 2 concentration in the chamber was measured using an infrared gas analyser (PP Systems Soil Respiration system, Stotfold, Hitchin, UK) for 120 s, and CO 2 flux calculated (in units of g CO 2 /m 2.h) from the rate of increase in CO 2.
0	Emissions of CO 2 were measured for all plots except those with surface straw, where the thick layer of straw made insertion of the chambers impractical.
0	Measurements were usually made in the mid-morning to early afternoon period.
0	To avoid compaction of soil in the vicinity of the plots during CO 2 measurements, wooden walkways were used to access the plots.
0	Carbon dioxide fluxes were monitored as described above between 7 April and 18 June 2001.
2	Towards the end of that period, the increase in chamber CO 2 concentration during the 120-s measurement period was quite small (generally <10 mL/L) and the system was approaching its detection limit.
0	Subsequent monitoring was done using chambers containing alkali traps to absorb emitted CO 2.
0	These chambers, which enclosed an area of 225 cm 2 , usually remained in place for 24 h, during which evolved CO 2 was absorbed in traps containing 20 mL of 0.25 M NaOH.
0	The absorbed CO 2 was determined by back titrating with 0.25 M HCl.
2	Comparisons carried out on 29 May and 7 June showed that CO 2 fluxes measured by the 2 techniques were very similar.
0	The final measurement using the alkali traps was on 30 August.
0	Cumulative CO 2 -C emissions were estimated from initiation of the experiment to 18 June (period during which the infrared analyser was used) and from then until the end of the experiment (alkali trap period) by trapezoidal integration beneath the CO 2 flux v. time curve.
0	The experiment was terminated after 158 days (10 September 2001).
0	Straw remaining on the surface straw plots was carefully collected so as to minimise soil contamination.
0	Soil was collected from each microplot in 50-mm increments (0-0.05, 0.05-0.10, 0.10-0.15, and 0.15-0.20 m depths).
0	The litter bags were retrieved and placed in plastic bags for transport to the laboratory.
0	Straw was taken from the litter bags, shaken gently on a 1-mm sieve to remove adhering soil, and weighed.
0	Samples were dried at 60 ˚C to determine moisture content.
0	Subsamples of dry, ground residue were heated in a muffle furnace at 550 ˚C for 5 h to determine ash content.
0	Further subsamples were taken to measure total C and N using a LECO C/N/S analyzer (LECO Corporation, St Louis, MI).
0	Residue mass and residue C and N contents were corrected to an ash-free basis to account for variable amounts of contaminating soil.
0	Surface straw was treated similarly to litter bag straw.
0	Soils from the soil-mixed straw treatment were oven-dried (30 ˚C) for at least 16 h, after which large straw pieces were removed by hand.
0	Further straw separation was achieved by passing the soil through a small, laboratory-based thresher fitted with a 5-mm sieve.
0	Finally, smaller pieces of straw were recovered using a seed dresser fitted with 4-and 2-mm screens.
0	Stones were removed from the separated residue by hand. Straw collected from the 4 sampled depths of each plot was combined.
0	Residues were recovered from soil taken from the control plots using a combination of hand picking (larger pieces), and sieving using the thresher and seed dresser, as outlined above.
0	The recovered residue was dried and analysed for ash, C, and N. Straw recovery from the soil-mixed straw treatment was estimated by subtracting the mean amount of ash-free residue measured in the 5 control plots.
0	Soil mineral N at end of experiment Mineral N was extracted from subsamples of field-moist soil from each depth increment (0-0.05, 0.05-0.10, 0.10-0.15, 0.15-0.20 m) by extraction with 2 M KCl and measuring nitrate-and ammonium-N using standard colourimetric procedures (Keeney and Nelson 1982).
0	The data were examined with analysis of variance (ANOVA) using GENSTAT 8 (GenStat 2005).
0	The CO 2 flux data for each measurement date were analysed separately.
0	A probability level of 0.05 was used to determine significance.
0	When straw was applied, the soil was moderately warm (mean daily temperature in the top 0.10 m ~14 ˚C), but dry (moisture content of 14% w/w).
0	Soil temperature remained relatively high (generally 12-14 ˚C in top 0.10 m) during the following 40 days, after which there was a rapid decline with the onset of winter (Fig. 1 ).
0	Soil temperature during the winter period of mid May to the end of August (days 41 to 147) was generally below about 6-7 ˚C in the 0-0.10 m layer (mean 5.4 ˚C).
0	Total rainfall during the experiment was 181 mm, which is less than the long-term mean for the period (~300 mm).
0	During the first 28 days, there was a total of only 0.6 mm of rain, and dry soil conditions persisted until 17 mm of rain fell between days 29 and 35.
1	Decomposition of low N (high C : N ratio) straws can be limited by N availability, particularly in the early phase of decomposition (Christensen 1986 ).
0	The straws used in this study had C : N ratios that were not especially high (64 for barley and 89 for wheat; Table 1 ) and soil mineral N at initiation of the trial was relatively high (117 kg/ha in the 0-0.20 m layer).
1	Thus, N availability is unlikely to have limited decomposition of incorporated straw, though it may be one of the factors limiting decomposition of surface straw (Wang et al. 2001).
2	Despite the relatively low soil moisture content, there was a large pulse of CO 2 production in the first 4 days where straw was mixed through the soil (Fig. 2 ).
2	Litter bag straw resulted in smaller increases in CO 2 emissions during this initial period.
2	Emissions of CO 2 from the soil-mixed straw plots decreased sharply following the initial pulse, though they remained above those measured in litter bag plots.
2	There was a trend for CO 2 emissions from all treatments to decrease from about day 8 to ~day 24 (30 April).
2	Measurements made on day 28 (4 May), before the 17 mm of rain between days 29 and 35, showed that the soil-mixed straw treatment continued to have a significantly (P < 0.05) higher output of CO 2 than the litter bag treatment (mean of 0.17 v. 0.12 g/m 2.h).
2	In response to rainfall on days 29-35, there was another pulse of CO 2 production with a peak at about day 33.
2	During this period, values for soil-mixed straw were again substantially higher than for the litter bags.
0	By day 40 (mid May), the soil was approaching field capacity (Table 2 ) and temperature was beginning to decline (Fig. 1 ).
1	Low temperature was likely the major abiotic limitation to straw decomposition during the remainder of the experiment.
2	From about day 55 (late May) through to the end of the experiment, CO 2 fluxes were generally higher for the straw-treated plots than from the control, but effects of straw placement were small or non-significant.
1	The high initial CO 2 output where straw was mixed through the soil was likely due to rapid decomposition of labile compounds in straw, as observed in previous work (Cogle et al. 1989).
2	Even though the soil was relatively dry, soil-mixed straw apparently absorbed sufficient moisture through extensive soil-straw contacts to enable microorganisms to rapidly initiate decomposition of labile components such as the water-soluble material, which made up 7-9% of straw-C (Table 1 ).
1	Fungi, which are more tolerant of low soil water potentials than are bacteria (Paul and Clark 1989) , are likely to have played a dominant role during this early decomposition stage.
1	Limited soil-straw contact in the case of the litter bags reduced the decomposition, probably by slowing absorption of water from the soil and delaying microbial colonisation of the straw.
0	In this field experiment, effects of within-soil straw placement persisted over a relatively long period (~2 months from time of incorporation).
2	This contrasts with observations from laboratory studies where decomposition rate of localised straw was slower than that of soil-mixed straw during the initial 7 days only (Parr and Reuszer 1959).
2	Total output of CO 2 -C from initiation of the experiment to 18 June (period during which the infrared gas analyser was used to measure CO 2 emissions), averaged 83 g/m 2 for soil-mixed straw plots, 61 g/m 2 for the litter bag plots, and 34 g/m 2 in the nostraw control (Table 3 ).
2	Amounts of CO 2 -C released during the remainder of the experiment were little affected by straw placement.
2	The incorporated straw accounted for a large part of the emitted CO 2 -C, particularly where straw was mixed through the soil (Table 3 ).
0	Soil organic matter had been depleted due to a history of intensive cropping, and as a result, background soil respiration was relatively low.
2	The CO 2 -C attributable to straw (estimated by subtracting CO 2 -C emitted from no-straw control from that released from the straw treated plots) was almost twice as high for the soil-mixed straw v. litter bag straw (averaged over straw type) until 18 June (first 73 days).
2	Over the entire experiment, CO 2 -C generated from soilmixed straw was, on average, 1.66 times that from litter bag straw (Table 3 ).
1	These results demonstrate that within-soil placement can have a strong and persistent influence on straw decomposition.
1	Results obtained using the litter bag technique may seriously underestimate decomposition rate of straw when it is incorporated using an implement that distributes it through the soil matrix.
1	The importance of straw placement/distribution is likely to be greatest when the soil and straw are dry at the time of incorporation.
1	Under moist conditions, localised straw would wet up relatively rapidly and the decomposition process may be less affected by within-soil straw placement.
0	However, in many grain-producing areas, soil can be depleted of available moisture by harvest, and under these conditions, mixing straw through the soil may be a practical way to speed up decomposition prior to establishing the next crop.
1	Straw type did not have a strong influence on CO 2 emissions.
2	Estimated straw-derived CO 2 -C (averaged over placement methods) was 55 g/m 2 for barley straw v. 47 g/m 2 for wheat straw (difference significant, P = 0.03).
1	There was no evidence of an interaction between straw type and incorporation method.
1	There is evidence in the literature that barley straw decomposes slightly more rapidly than wheat straw (Summerell and Burgess 1989) , although in a litter bag experiment, Christensen (1985) found no difference between wheat and barley straws.
1	The slightly greater decomposition of barley straw may be a reflection of its higher N and water-soluble C contents (Table 1 ).
2	Straw recovered from the soil at the end of the experiment was substantially less when it had been soil-mixed than when it was enclosed in litter bags (Table 4 ).
2	Mass loss of straw was 31-33% for the litter bags, whereas 61-71% of the soil-mixed straw was not recovered by careful hand picking and mechanical sieving.
2	As expected, decomposition of surface straw was much less than that of incorporated straw; mass loss of surface straw was <15% for both barley and wheat straw (Table 4 ).
2	Residue mass loss data for the litter bags are consistent with other local measurements.
2	Beare et al. (2002) reported mass loss of barley straw from litter bags of about 35% in the period 5 May-15 September, similar to our values of 31-33% between 6 April and 10 September.
2	Our mass loss values for litter bag straw also compare well with predictions of the Douglas and Rickman (1992) model.
0	That model predicts straw mass loss based on air temperature (degree-days above a base of 0 ˚C), placement (incorporated or surface-placed straw), and initial straw N content.
1	An assumption implicit in the model is that incorporated residues decompose at the same rate regardless of whether they are localised or mixed through the soil.
1	Predicted mass loss by the Douglas-Rickman model for incorporated straw was 42% for barley and 39% for wheat.
2	Predicted loss for surface straw was 12-13%. These results suggest that that the Douglas-Rickman model is useful in predicting decomposition of surface-placed residues (no-tillage scenario) and residues that are localised within the soil.
1	However, the model may systematically underpredict decomposition of residues that are well distributed through the soil.
1	It is likely that mass loss of soil-mixed straw was overestimated to some extent because of the difficulty of recovering all of the undecomposed residue.
0	Tests of recovery efficiency showed that, after initial mixing of straw into microplots, essentially all of the straw was recovered using our extraction method (Francis et al. 1999).
1	However, the efficiency of recovery may decrease as decomposition progresses, because straw becomes brittle and more fragmented, making it difficult to extract.
0	From a practical perspective, straw fragments not recovered by our extraction technique would pose no difficulties for the farmer as they would not interfere with mechanical operations such as seedbed preparation and seed drilling.
2	As would be expected, straw C recovery values were very similar to those for straw mass (ash-free basis) recovery (Table 4 ).
1	There was evidence of a straw type  placement interaction on straw mass and C recovery; recovery values for barley were significantly less than those of wheat straw in the soil-mixed treatment.
1	Straw type had a strong influence on straw N recovery.
2	On average, recovery of wheat straw N was 70%, compared with 57% for barley straw.
2	The barley contained a significant amount of nitrate-N (640 mg N/kg v. 16 mg N/kg for wheat straw: Table 1 ).
1	Leaching of NO 3 -N, which made up almost 10% of total N in barley straw, would largely explain the lower recovery of N for barley straw.
1	The surface straw lost relatively more N than C. While leaching of NO 3 -N may partly account for the N loss from barley straw, leaching of soluble organic N must also have occurred to account for N losses from surface straw.
1	Low N recovery from soil-mixed straw is likely to be partly due to failure to extract all of the undecomposed straw fragments.
1	Residues with high C : N ratios can cause temporary immobilisation of soil N after incorporation (Parr and Papendick 1978).
1	Soil mineral N levels (0-0.20 m) at the end of the experiment were little affected by the straw treatments (Table 2 ).
1	The final mineral N levels were much lower than at the beginning of the experiment (~20 v. 117 kg/ha), indicating that there was significant loss of N during the winter period, presumably by leaching and denitrification.
1	Any straw-induced N immobilisation would not be detectable, as leaching/denitrification had an over-riding influence on mineral N changes during the experiment.
0	Evolved CO 2 -C v. C recovery
2	Straw recovery data are consistent with the CO 2 -C output values, in that both approaches suggest that decomposition was less in litter bags than when straw was mixed into the soil.
2	However, the estimates of straw decomposition based on CO 2 -C output were substantially less than those obtained by mass loss.
2	For example, mass loss of C from the litter bags averaged 96 g/m 2 , compared with 38 g/m 2 of straw-C respired as CO 2.
0	Explanations might include either underestimation of CO 2 -C output or incomplete recovery of undecomposed straw.
0	Fluxes of CO 2 -C were monitored frequently during the high emission period following straw addition, with less frequent measurements during the cool winter period (Fig. 2 ); it is unlikely that any high CO 2 emission episodes were missed using this sampling protocol.
0	Measurements using the infrared gas analyser (first 73 days) were usually made in mid-morning/early afternoon when temperature was close to the diurnal maximum.
0	As no adjustment was made for temperature-related diurnal variation in CO 2 emissions (Akinremi et al. 1999) , it is likely that CO 2 -C output was overestimated rather than underestimated.
2	Measurements of CO 2 emissions ceased about 10 days before straw recovery; however, estimates obtained by extrapolating the data in Fig.
2	2 suggest that straw-C respired as CO 2 during that 10-day period would be quite small (<2 g C/m 2 ).
2	Cumulative CO 2 -C evolution represented an average of 21% of added C where straw was soil-mixed.
2	This does not appear unrealistically low compared with some reports.
2	For example, in a year-long study, Jensen et al. (1997) found that only 23-31% of residue C (residue incorporated by rotovating) was respired as CO 2 , whereas mass loss would be expected to be ~70%.
2	In a laboratory experiment in which canola (Brassica napus) residues were mixed into soil, 33-35% of residue-C was respired in a 6-month incubation at 20 ˚C (Bhupinderpal-Singh et al. 2006).
0	The litter bag technique is a well-established method and there is no reason to believe that there was any significant loss of undecomposed straw during recovery of the bags from the plots, transport to the laboratory, or in subsequent processing.
1	There may have been some leaching of water-soluble organic matter from the litter bags during the winter, but leaching probably would not make a large contribution to straw mass loss (Berg and McClaugherty 2003).
0	Losses from the litter bags may have occurred due to fragmentation or removal by earthworms and arthropods, but the extent of such losses is difficult to quantify.
0	Very few earthworms were noted during set-up of the experiment, suggesting that straw removal by earthworms would be minimal.
1	Only part of the decomposed straw C is likely to have been respired as CO 2 ; some will also have been assimilated into microbial biomass and microbial metabolites (Paul and Clark 1989; Witter and Dahlin 1995; Devevre and Horwath 2000).
0	The relative amounts of straw C respired as CO 2 -C v. synthesised into microbial tissue depends on microbial growth efficiency, which is calculated from the equation:
2	Utilisation efficiency values used in soil C turnover models are: 60% for the easily decomposable straw fraction (~15% of straw mass); 40% for the slowly decomposable straw fraction; and 10% for lignin (Paul and Clark 1989).
1	These microbial efficiency values are consistent with recent measurements that indicated y was in the range 40-60% for rice straw during a 160-day decomposition period (Devevre and Horwath 2000).
2	Substituting mean values for litter bag straw (i.e. C s = 96 g C/m 2 and C m = 38 g C/m 2 ) into Eqn 1 gives a value of 60% for y, which is at the high end of the range observed by Devevre and Horwath (2000) and similar to the model value for the easily decomposable straw (Paul and Clark 1989).
2	 where C s is the total C decomposed, C m is C respired as CO 2 , and y is the efficiency of C use for biosynthesis of new biomass (Paul and Clark 1989).
1	Decomposing straw is known to support a large microbial population (Beare et al. 2002) but that microbial C would be measured as part of the recovered straw C. For microbial C synthesis to explain the difference between C loss from litter bags and CO 2 -C output, the microbial C would have to be located outside the litter bag so as not to be measured as part of the recovered straw C. It has been demonstrated that microbial biomass in soil above and below buried litter bags may contain substantial quantities of residue-derived C (Fliessbach et al. 1995).
1	Recently, evidence has been presented that filamentous fungi can translocate straw C into the soil while transferring N in the opposite direction (Frey et al. 2003).
1	The latter workers studied fungally mediated C translocation from surface straw; however, this mechanism is likely to operate to an even greater extent with buried bag straw because the soil-straw interface is larger (above and below the straw layer).
2	Frey et al. (2003) observed that, where 13 C-labelled wheat straw (~2 t/ha) was placed on the soil surface for 35 days, significant straw-derived C (120-150 mg C/kg soil) was measured in the underlying 25 mm soil layer as a result of fungal translocation.
1	Extrapolation of these findings to our study suggests that C translocation by fungi to the soil above and below the litter bags could amount to 20-30 g C/m 2 , assuming that translocated C increases in proportion to the straw application rate.
0	Unfortunately, no measurements were made to quantify changes in soil microbial biomass.
0	Although fungal C translocation may offer a partial explanation for the difference between mass loss of C and CO 2 -C evolved from the litter bags, further work is needed to examine the role of fungal C translocation as a mechanism of C loss from buried straw.
2	Evolved CO 2 -C accounted for an average of about 40% of straw C lost from the litter bags, whereas about 30% of the unrecovered C in the soil-mixed straw treatments could be attributed to C loss as CO 2 (Table 4 ).
0	This would imply that straw C recovery was not greatly underestimated in the soilmixed treatments.
2	Based on the ratio of CO 2 -C evolved to C mass loss observed for the litter bags, loss of straw C in the straw-mixed treatments would average 160 g/m 2 , compared with the measured value of 209 g/m 2.
1	The difference may be assumed to represent C in straw fragments that were too small to be physically recovered from the soil.
0	These small, partially decomposed straw fragments could legitimately be considered part of the soil organic matter, most probably the light fraction or particulate organic matter.
1	Increases in light fraction organic matter (material recoverable by flotation on a dense liquid) following incorporation of straw have been demonstrated by Magid et al. (1997b).
0	A strict definition of decomposition is difficult and residue decomposition has, by default, been defined by the method used to measure it (Berg and McClaugherty 2003).
0	Residue mass loss comprises C respired as CO 2 , C partitioned to the microbial biomass, and C in residue fragments that become part of soil C pools such as light fraction and particulate organic matter.
0	The latter may represent a relatively large part of the mass loss where unconfined straw is used.
0	Models capable of predicting decomposition of unconfined straw could assist greatly in developing practical management practices for post-harvest residues.
0	For this purpose, straw that has physically fragmented beyond a certain point (e.g. passes through a 2-mm sieve) could effectively be regarded as decomposed, although that criterion may not fit with some definitions of decomposition.
0	Where these residue fragments need to be specifically quantified, this may be achieved using density separation techniques (Magid et al. 1997a).
1	This study has confirmed that straw mixed through the soil decomposes more rapidly than when it is localised, as occurs when plough-incorporated.
0	Effects of within-soil placement on decomposition persisted for about 2 months following incorporation.
2	Over the course of the experiment, cumulative CO 2 evolution from soil-mixed straw was 1.7 times that from litter bag straw.
2	Straw mass recovered at the end of the experiment averaged 68% for litter bag straw compared with 34% for soil-mixed straw and 86% for surface-placed straw.
2	Carbon respired as CO 2 -C accounted for 29-46% of mass loss of incorporated straw-C. The unrecovered straw-C presumably accumulated in soil C pools such as microbial biomass and particulate organic matter, with the latter likely to be an important repository where straw was mixed through the soil.
1	Straw type had only small effects CO 2 -C emissions and mass loss, although there was evidence that barley straw decomposed more rapidly than wheat straw, possibly because it was higher in water-soluble C and had a higher N content than wheat straw.
1	The litter bag technique may seriously underestimate decomposition rate of straw that is mixed through the soil, and decomposition models need to be revised to take within-soil placement into account.